Tritrophic interactions have a central role in ecological theory and applications [1-3]. Particularly, systems comprised of plants, herbivores and predators have historically received wide attention given their ubiquity and economic importance [4]. Although ecologists have long aimed to understand the forces that govern alternating ecological effects at successive trophic levels [5], several key open questions remain (at least partially) unanswered [6]. In particular, the analysis of complex food webs has questioned whether ecosystems can be viewed as a series of trophic chains [7,8]. Moreover, whether systems are mostly controlled by top-down (trophic cascades) or bottom-up processes remains an open question [6].
Traditionally, studies have addressed how species diversity at different food chain compartments affect the strength and direction of trophic cascades [9]. For example, many studies have tested whether biological control was more efficient with more than one species of natural enemies [10-12]. Much less attention has been given to the role of within-species variation in shaping trophic cascades [13]. In particular, whereas the impact of trait variation within species of plants or predators on successive trophic levels has been recently addressed [14,15], the impact of intraspecific herbivore variation is in its infancy (but see [16]). This is at odds with the resurgent acknowledgment of the importance of individual variation for several ecological processes operating at higher levels of biological organization [17].
Sources of variation within species can come in many flavours. In herbivores, striking ecological variation can be found among populations occurring on different host plants, which become genetically differentiated, thus forming host races [18,19]. Curiously, the impact of variation across host races on the strength of trophic cascades has, to date, not been explored. This is the gap that the manuscript by Sentis and colleagues [20] fills. They experimentally studied a curious tri-trophic system where the primary consumer, pea aphids, specializes in different plant hosts, creating intraspecific variation across biotypes. Interestingly, there is also ecological variation across lineages from the same biotype. The authors set up experimental food chains, where pea aphids from different lineages and biotypes were placed in their universal legume host (broad bean plants) and then exposed to a voracious but charming predator, ladybugs. The full factorial design of this experiment allowed the authors to measure vertical effects of intraspecific variation in herbivores on both plant productivity (top-down) and predator individual growth (bottom-up).
The results nicely uncover the mechanisms by which intraspecific differences in herbivores precipitates vertical modulation in food chains. Herbivore lineage and host-plant specialization shaped the strength of trophic cascades, but curiously these effects were not modulated by density-dependence. Further, ladybugs consuming pea aphids from different lineages and biotypes grew at distinct rates, revealing bottom-up effects of intraspecific variation in herbivores.
These findings are novel and exciting for several reasons. First, they show how intraspecific variation in intermediate food chain compartments can simultaneously reverberate both top-down and bottom-up effects. Second, they bring an evolutionary facet to the understanding of trophic cascades, providing valuable insights on how genetically differentiated populations play particular ecological roles in food webs. Finally, Sentis and colleagues’ findings [20] have critical implications well beyond their study systems. From an applied perspective, they provide an evident instance on how consumers’ evolutionary specialization matters for their role in ecosystems processes (e.g. plant biomass production, predator conversion rate), which has key consequences for biological control initiatives and invasive species management. From a conceptual standpoint, their results ignite the still neglected value of intraspecific variation (driven by evolution) in modulating the functioning of food webs, which is a promising avenue for future theoretical and empirical studies.

References

[1] Price, P. W., Bouton, C. E., Gross, P., McPheron, B. A., Thompson, J. N., & Weis, A. E. (1980). Interactions among three trophic levels: influence of plants on interactions between insect herbivores and natural enemies. Annual review of Ecology and Systematics, 11(1), 41-65. doi: 10.1146/annurev.es.11.110180.000353
[2] Olff, H., Brown, V.K. & Drent, R.H. (1999). Herbivores: between plants and predators. Blackwell Science, Oxford.
[3] Tscharntke, T. & Hawkins, B.A. (2002). Multitrophic level interactions. Cambridge University Press. doi: 10.1017/CBO9780511542190
[4] Agrawal, A. A. (2000). Mechanisms, ecological consequences and agricultural implications of tri-trophic interactions. Current opinion in plant biology, 3(4), 329-335. doi: 10.1016/S1369-5266(00)00089-3
[5] Pace, M. L., Cole, J. J., Carpenter, S. R., & Kitchell, J. F. (1999). Trophic cascades revealed in diverse ecosystems. Trends in ecology & evolution, 14(12), 483-488. doi: 10.1016/S0169-5347(99)01723-1
[6] Abdala‐Roberts, L., Puentes, A., Finke, D. L., Marquis, R. J., Montserrat, M., Poelman, E. H., ... & Mooney, K. (2019). Tri‐trophic interactions: bridging species, communities and ecosystems. Ecology letters, 22(12), 2151-2167. doi: 10.1111/ele.13392
[7] Polis, G.A. & Winemiller, K.O. (1996). Food webs. Integration of patterns and dynamics. Chapmann & Hall, New York. doi: 10.1007/978-1-4615-7007-3
[8] Torres‐Campos, I., Magalhães, S., Moya‐Laraño, J., & Montserrat, M. (2020). The return of the trophic chain: Fundamental vs. realized interactions in a simple arthropod food web. Functional Ecology, 34(2), 521-533. doi: 10.1111/1365-2435.13470
[9] Polis, G. A., Sears, A. L., Huxel, G. R., Strong, D. R., & Maron, J. (2000). When is a trophic cascade a trophic cascade?. Trends in Ecology & Evolution, 15(11), 473-475. doi: 10.1016/S0169-5347(00)01971-6
[10] Sih, A., Englund, G., & Wooster, D. (1998). Emergent impacts of multiple predators on prey. Trends in ecology & evolution, 13(9), 350-355. doi: 10.1016/S0169-5347(98)01437-2
[11] Diehl, E., Sereda, E., Wolters, V., & Birkhofer, K. (2013). Effects of predator specialization, host plant and climate on biological control of aphids by natural enemies: a meta‐analysis. Journal of Applied Ecology, 50(1), 262-270. doi: 10.1111/1365-2664.12032
[12] Snyder, W. E. (2019). Give predators a complement: conserving natural enemy biodiversity to improve biocontrol. Biological control, 135, 73-82. doi: 10.1016/j.biocontrol.2019.04.017
[13] Des Roches, S., Post, D. M., Turley, N. E., Bailey, J. K., Hendry, A. P., Kinnison, M. T., ... & Palkovacs, E. P. (2018). The ecological importance of intraspecific variation. Nature Ecology & Evolution, 2(1), 57-64. doi: 10.1038/s41559-017-0402-5
[14] Bustos‐Segura, C., Poelman, E. H., Reichelt, M., Gershenzon, J., & Gols, R. (2017). Intraspecific chemical diversity among neighbouring plants correlates positively with plant size and herbivore load but negatively with herbivore damage. Ecology Letters, 20(1), 87-97. doi: 10.1111/ele.12713
[15] Start, D., & Gilbert, B. (2017). Predator personality structures prey communities and trophic cascades. Ecology letters, 20(3), 366-374. doi: 10.1111/ele.12735
[16] Turcotte, M. M., Reznick, D. N., & Daniel Hare, J. (2013). Experimental test of an eco-evolutionary dynamic feedback loop between evolution and population density in the green peach aphid. The American Naturalist, 181(S1), S46-S57. doi: 10.1086/668078
[17] Bolnick, D. I., Amarasekare, P., Araújo, M. S., Bürger, R., Levine, J. M., Novak, M., ... & Vasseur, D. A. (2011). Why intraspecific trait variation matters in community ecology. Trends in ecology & evolution, 26(4), 183-192. doi: 10.1016/j.tree.2011.01.009
[18] Drès, M., & Mallet, J. (2002). Host races in plant–feeding insects and their importance in sympatric speciation. Philosophical Transactions of the Royal Society of London. Series B: Biological Sciences, 357(1420), 471-492. doi: 10.1098/rstb.2002.1059
[19] Magalhães, S., Forbes, M. R., Skoracka, A., Osakabe, M., Chevillon, C., & McCoy, K. D. (2007). Host race formation in the Acari. Experimental and Applied Acarology, 42(4), 225-238. doi: 10.1007/s10493-007-9091-0
[20] Sentis, A., Bertram, R., Dardenne, N., Simon, J.-C., Magro, A., Pujol, B., Danchin, E. and J.-L. Hemptinne (2020) Intraspecific difference among herbivore lineages and their host-plant specialization drive the strength of trophic cascades. bioRxiv, 722140, ver. 4 peer-reviewed and recommended by PCI Ecology. doi: 10.1101/722140

Across the temporal expanse of history, the impact of human activities on global landscapes has manifested as a complex interplay of ecological alterations. From the advent of early agricultural practices to the successive waves of industrialization characterizing the 18th and 19th centuries, anthropogenic forces have exerted profound and enduring transformations upon Earth's ecosystems. Indeed, by 2017, more than 80% of the terrestrial biosphere was transformed by human populations and land use, and just 19% remains as wildlands (Ellis et al. 2021).
 
Urbanization engenders profound alterations in environmental conditions, exerting substantial impacts on biological communities. The expansion of built infrastructure, modification of land use patterns, and the introduction of impervious surfaces and habitat fragmentation are key facets of urbanization (Faeth et al. 2011). These alterations generate biodiversity loss, changes in the composition of biological communities, disruptions in access and availability of food and nutrients, and a loss of efficiency in the immune system's control of infections, etc. (Reyes et al. 2013).
 
In this study, Caizergues et al. (2023) investigated the prevalence and diversity of avian malaria parasites (Plasmodium/Haemoproteus sp. and Leucocytozoon sp.) in great tits (Parus major) living across an urbanization gradient. The study reveals nuanced patterns of avian malaria prevalence and lineage diversity in great tits across urban and non-urban environments. While overall parasite diversity remains consistent, there are marked differences in prevalence between life stages and habitats. They observed a high prevalence in adult birds (from 95% to 100%), yet lower prevalence in fledglings (from 0% to 38%). Notably, urban nestlings exhibit higher parasite prevalence than their non-urban counterparts, suggesting a potential link between early malaria infection and the urban heat island effect. This finding underscores the importance of considering both spatial and temporal aspects of urbanization in understanding disease dynamics. Parasite lineages were not habitat-specific. The results suggest a potential parasitic burden in more urbanized areas, with a marginal but notable effect of nest-level urbanization on Plasmodium prevalence. This challenges the common perception of lower parasitic prevalence in urban environments and highlights the need for further investigation into the factors influencing parasite prevalence at finer spatial scales.
 
The discussion emphasizes the significance of examining vector distributions, abundance, and diversity in urban areas, which may be influenced by ecological niches and the presence of suitable habitats such as marshes. The identification of habitat-specific Haemosporidian lineages, particularly those occurring more frequently in urban areas, raises intriguing questions about the factors influencing parasite diversity. The presence of rare lineages in urban environments, such as AFR065, DELURB4, and YWT4, suggests a potential connection between urban bird communities and specific parasite strains.
 
Future research should empirically demonstrate these relationships to enhance our understanding of urban parasitology. This finding has broader implications for wildlife epidemiology, especially when introducing or keeping exotic wildlife in contact with native species. The study highlights the importance of considering not only the prevalence but also the specific lineages of parasites in understanding the dynamics of avian malaria in urban and non-urban habitats. This preprint contributes valuable insights to the ongoing discourse on the intricate interplay between ecological repercussions of human-induced changes (urbanization), biological communities, and the prevalence of vector-borne diseases.
 
References

Caizergues AE, Robira B, Perrier C, Jeanneau M, Berthomieu A, Perret S, Gandon S, Charmantier A (2023) Cities as parasitic amplifiers? Malaria prevalence and diversity in great tits along an urbanization gradient. bioRxiv, 2023.05.03.539263, ver. 3 peer-reviewed and recommended by Peer Community in Ecology. https://doi.org/10.1101/2023.05.03.539263

Ellis EC, Gauthier N, Klein Goldewijk K, Bliege Bird R, Boivin N, Díaz S, Fuller DQ, Gill JL, Kaplan JO, Kingston N, Locke H, McMichael CNH, Ranco D, Rick TC, Shaw MR, Stephens L, Svenning JC, Watson JEM. People have shaped most of terrestrial nature for at least 12,000 years. Proc Natl Acad Sci U S A. 2021 Apr 27;118(17):e2023483118. https://doi.org/10.1073/pnas.2023483118. 

Faeth  SH, Bang  C, Saari  S (2011) Urban biodiversity: Patterns and mechanisms. Ann N Y Acad Sci 1223:69–81. https://doi.org/10.1111/j.1749-6632.2010.05925.x

Faeth  SH, Bang  C, Saari  S (2011) Urban biodiversity: Patterns and mechanisms. Ann N Y Acad Sci 1223:69–81. https://doi.org/10.1111/j.1749-6632.2010.05925.x

Reyes  R, Ahn  R, Thurber  K, Burke  TF (2013) Urbanization and Infectious Diseases: General Principles, Historical Perspectives, and Contemporary Challenges. Challenges Infect Dis 123. https://doi.org/10.1007/978-1-4614-4496-1_4

Disentangling the effects on ecosystem structure and functioning of natural and human-induced impacts in transitional waters is a great challenge in coast ecology. This is due to the observation that the ecosystems of transitional waters are naturally dynamic systems with characteristics of stressed systems. For example, the benthic communities present low species richness and high abundance of species with a high tolerance to variations, e.g., salinity. This general observation is known as the paradigm of the “Transitional Waters Quality Paradox” (Zaldívar et al., 2008) derived from the previously described “Estuarine Quality Paradox” (Elliott and Quintino, 2007). 

In Jones et al. (2024) “Disentangling the effects of eutrophication and natural variability on macrobenthic communities across French coastal lagoons”, a great diversity of lagoons is analyzed to disentangle the effects of eutrophication from those of natural environmental variability on benthic macroinvertebrates and understanding the links between environmental variables affecting benthic macroinvertebrates. These authors use a very elegant set of numerical approaches, including correlograms, linear models and variance partitioning. They apply this suite to a dataset of macrobenthic invertebrate abundances and environmental variables from 29 Mediterranean coastal lagoons in France.

Through this suite of analyses, they demonstrate the strong complexity of the mechanisms interplaying in a situation of eutrophication on lagoon macrobenthos. The mechanisms involved are direct, like toxicity, or indirect, for example, through modifications of the sediment's biogeochemistry. Such a result on the different interactions involved is very important in the context of the search for indicators to define ecosystem status. Improving the definition of metrics is essential in environmental management decisions.

References

Elliott, M. and Quintino, V. (2007) The estuarine quality paradox, environmental homeostasis and the difficulty of detecting anthropogenic stress in naturally stressed areas. Marine Pollution Bulletin 54, 640–645. https://doi.org/10.1016/j.marpolbul.2007.02.003

Jones et al. (2024) Disentangling the effects of eutrophication and natural variability on macrobenthic communities across French coastal lagoons bioRxiv, 2022.08.18.504439, ver. 4 peer-reviewed and recommended by Peer Community in Ecology. https://doi.org/10.1101/2022.08.18.504439

Zaldívar, J. (2008). Eutrophication in transitional waters: an overview. https://doi.org/10.1285/I18252273V2N1P1

​​​​​The persistence of populations facing the destruction of their habitat is a multifaceted question that has mobilized theoreticians and empiricists alike for decades. As an ecological question, persistence has been studied as the spatial rescue of populations via dispersal into remaining suitable habitats. The spatial aggregation of habitat destruction has been a key component of these studies, and it has been applied to the problem of coexistence by integrating competition-colonization tradeoffs. There is a rich ecological literature on this topic, both from theoretical and field studies (Fahrig 2003). The relationship between life-history strategies of species and their resilience to spatially structured habitat fragmentation is also an important component of conservation strategies through the management of land use, networks of protected areas, and the creation of corridors. In the context of environmental change, the ability of species to adapt to changes in landscape configuration and availability can be treated as an eco-evolutionary process by considering the possibility of evolutionary rescue (Heino and Hanski 2001; Bell 2017). However, eco-evolutionary dynamics considering spatially structured changes in landscapes and life-history tradeoffs remains an outstanding question. Finand et al. (2023) formulate the problem of persistence in fragmented landscapes over evolutionary time scales by studying models for the evolution of dispersal in relation to habitat fragmentation and spatial aggregation. Their simulations were conducted on a spatial grid where individuals can colonize suitable patch as a function of their competitive rank that decreases as a function of their (ii) dispersal distance trait. Simulations were run under fixed habitat fragmentation (proportion of unsuitable habitat) and aggregation, and with an explicit rate of habitat destruction to study evolutionary rescue.

Their results reveal a balance between the selection for high dispersal under increasing habitat fragmentation and selection for lower dispersal in response to habitat aggregation. This balance leads to the coexistence of polymorphic dispersal strategies in highly aggregated landscapes with low fragmentation where high dispersers inhabit aggregated habitats while low dispersers are found in isolated habitats. The authors then integrate the spatial rescue mechanism to the problem of evolutionary rescue in response to temporally increasing fragmentation. There they show how rapid evolution allows for evolutionary rescue through the evolution of high dispersal. They also show the limits to this evolutionary rescue to cases where both aggregation and fragmentation are not too high. Interestingly, habitat aggregation prevents evolutionary rescue by directly affecting the evolutionary potential of dispersal. The study is based on simple scenarios that ignore the complexity of relationships between dispersal, landscape properties, and species interactions. This simplicity is the strength of the study, revealing basic mechanisms that can now be tested against other life-history tradeoffs and species interactions. Finand et al. (2023) provide a novel foundation for the study of eco-evolutionary dynamics in metacommunities exposed to spatially structured habitat destruction. They point to important assumptions that must be made along the way, including the relationships between dispersal distance and fecundity (they assume a positive relationship), and the nature of life-history tradeoffs between dispersal rate and local competitive abilities. 

References

Bell, G. 2017. Evolutionary Rescue. Annual Review of Ecology, Evolution, and Systematics 48:605–627. https://doi.org/10.1146/annurev-ecolsys-110316-023011 
Fahrig, L. 2003. Effects of Habitat Fragmentation on Biodiversity. Annual Review of Ecology, Evolution, and Systematics 34:487–515. https://doi.org/10.2307/30033784 
Finand, B., T. Monnin, and N. Loeuille. 2023. Evolution of dispersal and the maintenance of fragmented metapopulations. bioRxiv, 2022.06.08.495260, ver. 3 peer-reviewed and recommended by Peer Community in Ecology. https://doi.org/10.1101/2022.06.08.495260 
Heino, M., and I. Hanski. 2001. Evolution of Migration Rate in a Spatially Realistic Metapopulation Model. The American Naturalist 157:495–511. https://doi.org/10.1086/319927

Islands may represent just a small fraction (6.67%) of the planet’s land but they host a disproportionate 20% of the world’s biodiversity. Yet islands are highly vulnerable to human-induced change. Out of all IUCN Red list species, almost half of them are found on islands (Russell and Kueffer, 2019) while from the approximately 800 known extinctions that have occurred since the European expansion around the world, 75% have occurred on islands (IUCN, 2017).
 
Vulnerability is defined as “the degree to which a system is likely to experience harm due to exposure to a hazard” (Fuessel, 2007). It is meaningful to express vulnerability let s say of a population or species to a specific threat, pressure, or stress (like for instance the highly studied species vulnerability to climate change (Pacifici et al., 2015). Vulnerability is typically made up of three components: exposure (the extent of stress or threat that the species encounters and is projected to encounter), sensitivity (the ability of a species to persist under a given stress or threat), adaptation (the ability of the species to adapt to changes in a given stress or threat). 
 
When thinking of these three components, it becomes quickly evident that island biodiversity should be “naturally” vulnerable to global change stress (Frankham et al., 2002). First, it is hard to escape for insular species compared to mainland ones meaning that they cannot avoid exposure. Second, insular species are highly sensitive to any stress and stochastic events given their high specialisation due to their endemism. Third, insular species are less likely to adapt to new threats due to their small population sizes and naturally fragmented distribution ranges that both decrease their genetic diversity (aka adaptation potential). Thus, estimating the vulnerability of insular species is an important step towards better management and mitigation of their risk to extinction to ongoing global change. But an assessment framework designed for insular species is currently lacking.
 
Bellard and colleagues (Bellard et al., 2025) contribution is exactly addressing this objective. The authors present an adapted framework aimed to quantify the vulnerability of terrestrial insular biota by incorporating the idiosyncrasies of island biota: the island syndrome (ie the idiosyncratic evolutionary outcomes that arise in insular environments), the isolated nature of islands, and their high levels of endemism. It is the consequences of these three features that the authors highlight on expanding their insular vulnerability assessment. More in detail, Bellard et al (2025) build on existing vulnerability frameworks that are not specific to island ecosystems by focusing on the inclusion of multiple threats and enlarging the dimensions of diversity (taxonomic, functional and phylogenetic diversity). In that sense, this work stands out as it delivers a missing framework specific for island biodiversity, without minimising its potential as an extension on existing mainland (not island) vulnerability assessments.
 
The framework consists of 5 steps: 1) define the scope of the vulnerability assessment in terms of spatial and temporal extent, relevant threats, and studied biota; 2) determine the markers of exposure, sensitivity, and adaptive capacity; 3) compute measures of vulnerability and its components; and 4) conduct an uncertainty analysis to improve the vulnerability assessment. Step 5 is basically the use of the actual vulnerability assessment for practical conservation action and policy, and the authors are showing (Box 2 in Bellard et al (2025)) how their proposed vulnerability assessment could make the link to what is actually developed for (ie identifying which species are most vulnerable and what drives their vulnerability).
 
No doubt there is a growing number of literature on the design and application of biodiversity vulnerability assessments. Yet, this contribution is making the case for a special treatment of island biodiversity vulnerability assessments, while also providing a rather complete reading to a newcomer into vulnerability assessment frameworks.
 

References

Bellard Céline, Marino Clara, Butt Nathalie, Fernández-Palacios José María, Rigal François, Robuchon Marine, Lenoir Jonathan, Irl Severin, Benítez-López Ana, Capdevila Pol, Zhu G, Caetano Gabriel, Denelle Pierre, Philippe-Lesaffre Martin, Schipper Aafke, M Foden Wendy, Kissling W. Daniel, Leclerc Camille (2025) A framework to quantify the vulnerability of insular biota to global change. HAL, ver.3 peer-reviewed and recommended by PCI Ecology https://hal.science/hal-04550966

Frankham, R., Briscoe, D. A., and Ballou, J. D. (2002). Introduction to Conservation Genetics. Cambridge University Press.Fuessel, H.-M. (2007). Vulnerability: A generally applicable conceptual framework for climate change research. Glob. Environ. Change 17, 155–167. https://doi.org/10.1016/j.gloenvcha.2006.05.002

IUCN (2017). IUCN 2017 : International Union for Conservation of Nature annual report 2017. Available at: https://iucn.org/resources/annual-reports/iucn-2017-international-union-conservation-nature-annual-report-2017 (Accessed March 10, 2025).

Pacifici, M., Foden, W. B., Visconti, P., Watson, J. E. M., Butchart, S. H. M., Kovacs, K. M., et al. (2015). Assessing species vulnerability to climate change. Nat. Clim. Change 5, 215–224. https://doi.org/10.1038/nclimate2448

Russell, J. C., and Kueffer, C. (2019). Island Biodiversity in the Anthropocene. Annu. Rev. Environ. Resour. 44, 31–60. https://doi.org/10.1146/annurev-environ-101718-033245